16 Anaerobic Wastewater Treatment

Jules B. van Lier, Nidal Mahmoud and Grietje Zeeman

16.1 SUSTAINABILITY IN WASTEWATER TREATMENT

16.1.1 Definition and environmental benefits of anaerobic processes

The fermentation process in which organic material is degraded and biogas (composed of mainly methane and carbon dioxide) is produced, is referred to as anaerobic digestion. Anaerobic digestion processes occur in many places where organic material is available and redox potential is low (zero oxygen). This is typically the case in stomachs of ruminants, in marshes, sediments of lakes and ditches, municipal land fills , or even municipal sewers.

Anaerobic treatment itself is very effective in removing biodegradable organic compounds , leaving mineralised compounds like NH4-, PO/, S2- in the solution. Anaerobic treatment can be conducted in technically plain systems, and the process can be applied at any scale and at almost any place. Moreover the amount of excess sludge produced is very small and well stabilised, even having a market value when the so-called granular anaerobic sludge is produced in the bioreactor. Moreover, useful energy in the form of biogas is produced instead of high-grade energy

t 1

coommod. Aoooptiog thLbio dig,tioo io '"" merely removes organic pollutants, there are virtually few if any serious drawbacks left, even not with respect to the rate of start-up of the system. Figure 16.1 shows the fate of carbon and energy in both aerobic and anaerobic wastewater treatment (An WT) assuming that the oxidation of l kgCOD requires 1 kWh of aeration energy. In contrast to anaerobic treatment, aerobic treatment is genera lly characterised by high operational costs (energy), while a very large fract ion of the waste is converted to another type of waste (sludge). Aerobic treatment in a conventional activated sludge process yields about 50% (or more) new sludge from the COD converted , wh ich requires further treatment, e.g. anaerobic digestion, before it is reused, disposed off or incinerated. The carbon/energy flow principles of aerobic and anaerobic bio-conversion largely affect the set up of the corresponding wastewater treatment system . Not surprisingly, to date, An WT has evolved into a competitive wastewater treatment technology. Many different types of organically polluted wastewaters, even those that were previously believed not to be suitable for AnWT, are now treated by anaerobic high-rate conversion processes.

416

~ . HeoU"'' lnfl~ent B ----l,. Aeration -----r

(100 kWh) ~ Effluent , ' 12-10 kg COD

Influent

lurIRr+

Sludge, 30-60 kg

Biogas 40-45 m3 ( 70% CH4)

Anaerobic

Sludge, 5 kg Effluent

10-20 kg COD

Figure 16.1 Fate of carbon and energy in aerobic (above) and anaerobic (below) wastewater treatment

In countries like the Netherlands, almost all agro-industrial wastewaters are presently treated with anaerobic reactor systems. The application potential, e.g. in the petro-chemical industries , is rapidly growing. Figure 16.2 shows the gradual increase in the number of anaerobic high-rate reactors from the mid seventies onwards.

At present, a total number of 2,266 registered full scale installations are m operation, which are constructed by renowned companies like Paques, Biothane, Biotim, Enviroasia, ADI, Waterleau, Kurita,

2,400

2,200

2,000

1,800 ~ 1,600 c:

"' 0. 1, 400 ~ 0 1,200 a; .0 E 1,000 :::l z 800

600

400

200 0

(

Bio1ogical Wastewater ~ceatner c. pr ci~ t:S, Mode l 1ng Design

Degremont, Envirochemie, GWE, Grontmij as well as other local companies. To this number an estimated number of 500 'homemade' reactors can be added which are constructed by very small local companies or by the industries themselves but which do not appear in the statistics.

. Analysing the reasons why the selection for An WT was made, the following striking advantages of An WT over conventional aerobic treatment systems can be given:

reduction of excess sludge production up to 90%. up to 90% reduction in space requirement when

using expanded sludge bed systems. high applicable COD loading rates reaching 20-35 kg

COD per m3 of reactor per day, requiring smaller reactor volumes.

no use of fossil fuels for treatment, saving about I kWh/kgCOD removed, depending on aeration efficiency.

production of about 13.5 MJ CH4energy/kgCOD removed, giving 1.5 kWh electricity (assuming 40% electric conversion efficiency).

rapid start up ( < 1 week), using granular anaerobic sludge as seed material.

/. no or very little use of chemicals. plain technology with high treatment efficiencies. anaerobic sludge can be stored unfed, reactors can be

operated during agricultural campaigns only (e.g. 4 months per year in the sugar industry).

excess sludge has a market value. high rate systems facilitate water recycling m

factories (towards closed loops).

Nlf"'...cr--.ooo-o "' ,....... 1"-- r--. ,..... ,..... 00 0'0'0'0'-0'0'-0'-

~ N M ~ Lf"' ~ 1"-- 00 0' 0 ~ N M ~ Lf"' ...0 !"-- 00 0' 0 ~ N M ~ Lf"' ...0 00 00 00 00 00 00 00 00 00 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0 0 0 0 0 0 0 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0' 0 0 0 0 0 0 0

~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ N N N N N N N

Year

Figure 16.2 Increase in number of world wide installed anaerobic high-rate reactors in the period 1972-2006

Obviously, the exact ranking of the above advantages depends on the local economic and societal conditions. In the Netherlands, excess sludge handling is the cost-determining factor in operating wastewater treatment systems. Since land filling is no option for excess sewage sludge and biowastes, while prices for incineration reach 500/ton wet sludge or more, the low sludge production in anaerobic reactors is an immediate economic benefit. The system compactness, another important asset of An WT, can be illustrated by a full-scale example, where an anaerobic reactor with a 6 m diameter and a height of 25 m, suffices to treat up to 25 tons of COD daily. The produced sludge, which is less than 1 Ton dry matter per day in this example, is not a waste product, but is marketed as seed sludge for new reactors. Such compactness makes the system suitable for implementation on the industry premises or sometimes even inside the factory buildings. The latter is of particular interest in densely populated areas and for those industries aiming to use anaerobic treatment as the first step in a treatment for reclaiming process water.

The renewed interest in the energy aspects of An WT directly results from the ever rising energy prices and the overall concern on g,l'obal warming. The above 25 Tons COD/d of~?ustrial waste(water) can be converted in 7,000 m>CHi d (assuming 80% CH4 recovery), with an energy equivalent of about 250 GJ/d. Working with a modern combined heat power (CHP) gas engine, reaching 40% efficiency, a useful 1.2 MW electric power output can be achieved (Table 16.1). The overall energy recovery could even be higher (reaching up to 60%) if all the excess heat can be used on the industry premises or direct vicinity. Assuming that full aerobic treatment would require about 1 kWh/kgCOD removed, or 1 MW installed electric power in the above case, the total energy benefit of using An WT over the activated sludge process is 2.2 MW. At an energy price of 0.1 /kWh this equals about 5,000 /d. Apart from the energy itself, current drivers include the carbon credits that can be obtained by generating renewable energy using AnWT (Table 16.1). For an average coal-driven power plant, the generation of 1 MW electricity emits about 21 tonC02/d, whereas for a natural gas-

417

driven plant it is half that va lue. At a foreseen stabilised price of 20/ton C02, the above exampled industry could earn 500/d on carbon credits (based on a coal powered plant), whereas no fossil fuels are used for treating the wastewater. Although this amount is negligible in industrialised countries, it could provide a real incentive in developing countries to start treating the wastewater using high-rate An WT, and thereby protecting the local environment. The carbon credit policy can, therefore, be regarded as a Western subsidy for implementing An WT systems in less prosperous countries.

Table 16.1 gives a summary of the expected energy output as well as the predicted C02 emission reduction (if the produced CH4 is converted to electricity) of an anaerobic reactor, operated at commercially available organic loading rates .

16.2

16.2.1

MICROBIOLOGY OF ANAEROBIC CONVERSIONS

Anaerobic degradation of organic polymers

The anaerobic degradation pathway of organic matter is a multi step process of series and parallel reactions. This process of organic matter degradation proceeds in four success1ve stages, namely: (i) hydrolysis, (ii) acidogenesis, (iii) acetogenesis, and (iv) methanogenesis. These are discussed below.

Methanogenic bacteria are located at the end of the anaerobic food chain and, partly thanks to their activity, no large quantities of organic matter accumulate in anaerobic environments, where this matter is inaccessible to aerobic organisms. The anaerobic digestion process involves a complex food web, in which organic matter is sequentially degraded by a wide variety of micro-organisms. The microbial consortia involved jointly convert complex organic matter and ultimately mineralize it into methane (CH4), carbon dioxide C02, ammonium (NH3) , hydrogen sulphide (H2S) and water (H20).

Table 16.1 Energy output and CO, emission reduction applying anaerobic high-rate wastewater treatment systems

Loading capacity (kgCOD/m3.d) Energy output (MJ/m3 reactor installed per d) Electric power output (kW/m3 reactor installed) C02 emission reduct;on (tonCO/ m3 y, based on coal-driven power plant)

5 ~ 35 55 ~ 390 0.25 ~ 1.7 1.9 ~ 13

Assumptions: 80% CH, recovery relati ve to influent COD load and -+0% e lectric conversion efficiency using a modem combined heat power generator.

418

The anaerobic ecosystem is the result of complex interactions among microorganisms of several different species. The major groupings of bacteria and reaction they mediate are : (i) fermentative bacteria, (ii) hydrogen-producing acetogenic bacteria, (iii) hydrogen-consuming acetogenic bacteria, (iv) carbon dioxide-reducing methanogens, and (v) aceticlastic methanogens. The reactions they mediate are presented in Figure 16.3.

The digestion process may be subdivided into the following four phases:

Complex polymers

Hydrolysis 1

Amino acids, sugars

Fermentation

Acetate

Aceticlastic methanogenesis

Intermediary products (Propionate, Butyrate, etc.)

/

Anaerobic oxidation

Hydrogen Carbon dioxide

Homoacetogenesis

Methane Carbon dioxide

Hydrogenotrophic methanogenesis

Figure 16.3 Reactive scheme for the anaerobic digestion of polymeric materials. Numbers indicate the bacterial groups involved : 1. Hydrolytic and fermentative bacteria, 2. Acetogenic bacteria, 3 Homo-acetogenic bacteria, 4. Hydrogenotrophic methanogens, s. Aceticlastic methanogens (Gujer and Zehnder, 1983)

l) Hydrolysis, where enzymes excreted by fermentative bacteria (so-called ' exo-enzymes') convert complex, undissolved material into less complex, dissolved compounds which can pass through the cell walls and membranes of the fem1entative bacteria.

2) Acidogenesis, where the dissolved compounds present in cells of fermentative bacteria are converted into a number of simple compounds which are then excreted. The compounds produced during this phase include volatile fatty acids (VFAs), alcohols, lactic acid, C02, H2, NH3 and H2S, as well as new cell material.

3) Acetogenesis (intermediary acid production) where digestion products are convet1ed into acetate, hydrogen (H2) and C02, as well as new cell material.

4) Methanogenesis, where acetate, hydrogen plus carbonate, formate or methanol are converted into methane, C02 and new cell material.

In this global scheme, the following sub-processes can be distinguished (Figure 16.3):

I) Hydrolysis ofbiopolymers: hydrolysis of proteins hydrolysis of polysaccharides hydrolysis offats

2) Acidogenesis/fermentation: anaerobic oxidation of amino acids and sugars anaerobic oxidation of higher fatty acids and alcohols

3) Acetogenesis: formation of acetic acid and H2 from intermediary products (particularly VF As) homoacetogenesis: the formation of acetic acidfrom H2 and C02

4) Methanogenesis: methane formation from acetic acid methane formation from hydrogen and carbon dioxide

Figure 16.3 gives the unidirectional degradation of organic matter to the end products CH4 and C02. The homoacetogenic process illustrates the inter conversion of acetate, the major CH4 precursor and H2/C02 In practice, other back reactions may occur also, e.g. the formation of higher VFA or alcohols out of acetate and propionate. These back reactions are of particular importance in case of malfunctioning or perturbation of the anaerobic reactor or when a specific reaction is deliberately pursued. Under normal An WT applications, i.e. stable reactor performance under mesophilic conditions, acetate is the major precursor of CH4 (about 70% of the COD flux). Interesting to observe is that there is only COD conversion and no COD destruction. COD removal takes place owing to the fact that the end product of the reaction chain, CH4, is gaseous and highly insoluble in water.

In the case of the presence of alternative electron acceptors, like N03- and SO/, other bacterial groups will be present in the anaerobic reactor as well, such as denitrifiers and sulphate reducers (see Section 16.4).

16.2.1.1 Hydrolysis

Since bacteria are unable to take up particulate organic matter, the first step in anaerobic degradation consists of the hydrolysis of polymers . This process is merely a

'In . I , r

surface phenomena in which the polymeric particles are degraded through the action of exo-enzymes to produce smaller molecules which can cross the cell barrier, During the enzymatic hydrolys is process, proteins are hydrolyzed to amino acids, pol ysaccharide to simple sugars and lipids to long chain fatty acids (LCF A). Hydrolysis is in most cases, notably with (semi-) solid substrates and wastewaters with a high suspended solids (SS)/COD ratio , rate-limiting for the overall digestion process. Moreover, the hydrolysis process is very sensitive to temperature and temperature fluctuations. For that reason, the design of anaerobic digesters for (semi-) solid substrates and wastewaters with a high SS/COD ratio, such as distillery slops and low temperature sewage, is usually based on the hydrolysis step.

Hydrolysis can be defined as a process in which complex polymeric substrates, particulate or dissolved, are converted into monomeric and dimeric compounds which are readily accessible for the acidogenic bacteria. During anaerobic digestion of complex substrates hydrolysis is usually the first step. Although in some cases a preparatory step, i.e. physico-chemical pre-treatment or comminution, is needed to make hydrolysis possible. With the digestion of biological sludges, such as waste activated sludge, the hydrolysis of the sludge is preceded by death and lys is of the biomass. The hydrolysis is accomplished by exo-enzyms which are produced by the acidogenic bacteria. The products of the hydrolysis are the substrates for the acidogenic bacteria. A schematic presentation of the hydrolysis of lipids into LCF As is given in Figure 16.4.

0 11

0 CH20CR1 11 I

RzC-0-CH 0 I 11 CHz-0-CRJ

Triacylglycerol

0 R,.('

' o

CH20H 0 + 3 H20 Lipases HO CH + R2(' + 3 W

CH20H ' o

0 R3-C"

'o

Glycerol Long Chain Fatty Acids

Figure 16.4 The hydrolysis of lipids

As mentioned, hydrolysis is generally considered to be the rate-limiting step during the anaerobic digestion of complex substrates. However, usually thi s is not due to a lack of enzyme activity but to the avai labili ty of free accessible surface area of the particles and the overall structure of the solid substrate (Chandler et al .. 1980; Zeeman et al. , 1996). Even in dilute wastewaters such as low temperature domestic sewage, hydrolysi s

419

may determine the overall process and thereby determining the required reactor design. It must be noted that 45-75% of domestic sewage, and 80 % in primary sludge consists of suspended matter, The main biopolymers in sewage are proteins , carbohydrates and lipids.

16.2.1.2 Acidogenesis

During the acidogenesis step, the hydrolysis products (amino acids, simple sugars, LCFAs), which are relatively small soluble compounds, are diffused inside the bacterial cells through the cell membrane and subsequently fermented or anaerobically oxidized. Acidogenesis is a very common reaction and is performed by a large group of hydrolytic and non-hydrolytic microorganisms. About 1% of all known bacteria are (facultative) fermenters. The acidification products consist of a variety of small organic compounds, mainly VFAs, i.e. acetate and higher organic acids such as propionate and butyrate, as well as H2, C02, some lactic acids, ethanol and ammonia (Figure 16.3).

Characteristically, neutral compounds such as sugars and proteins are converted into VF As and carbonic acid, being the main end products. Therefore, fem1entative organisms are usually designated as acidifying or acidogenic microorganisms, and the process is therefore indicated by acidogenesis. Table 16.2 lists several acidogenic reactions starting from sucrose and generating different amounts of VF As, HC03-, H1, H+. Apparently, the type of end products depends on the conditions in the reactor medium. From Table 16.2 it follows that the L':.G 0 1 of the less energetic acidogenic reactions with sucrose as the substrate strongl y depends on the prevailing H2 concentrations. If H2 is effectively removed by H2 scavengmg organtsms such as methanogens, acetate will be the main end product. However, if methanogenesis is retarded and H2 accumulates, more reduced products such as propionate and butyrate are likely to appear and possibly the even more reduced compounds lactate and alcohols. Therefore, effluents of overloaded or perturbed anaerobic reactors (or reactors designed as acidifying reactors in an anaerobic two-step process) often contain these more reduced intermediate products.

Acidogenesis is the most rapid conversion step in the anaerobic food chai n. The L':.G 0 ' of acid ifying reactions is highest of all anaerobic conversions , resulting in ten to twentyfold higher bacterial growth rates, and tlvefold higher bacterial yields and conversion rates

4 20

Table 16.2 Acidogenic reactions w ith sucrose as the substrate and the corresponding free energy change (L1G 0 ') at 25C

Reactions L':.G 0 ' (kJ/mo1) Eq . C12H220 11 + 9H20 -f 4CH3COO- + 4HCOr + 8H+ + 8H2 -457.5 ( 16.1) C12H120 11 + 5H10 -f 2CH3CH2CH1COO- + 4HCO;- + 6H+ + 4H2 C12H220 11 + 3H20 -f 2CH3COO- + 2CH3CH1COO- + 2HCOs- + 6H+ + 2H2

- 554.1 - 610.5

(16.2) (16.3)

compared to methanogens (Table 16.3). For that reason, anaerobic reactors are subjected to souring, i.e. a sudden pH drop, when reactors are overloaded or perturbed by toxic compounds. Once alkalinity is consumed by the produced acids the pH starts to drop, resulting in a higher concentration of non-dissociated VF As, leading to a more severe inhibition of methanogens. The latter, obviously leads to an even quicker accumulation of VF As and subsequent pH drop (Figure 16.5).

Methane I Poor capacity buffering exceeded capacity

Methanogenic toxicity

VFA increases increasing

\ J pH decreases Unionized VFA

~ncreasing

Figure 16.5 Reactor pH drop as a result of methanogenic overloading and accumulating VFAs

The fact that acidifiers are active even at low pH ( 4), means the reactor souring to pH 4 to 5 can and will occur when the methanogenic capacity of the system is trespassed.

The acidogenic conversion of ami no acids generally follows the Stickland reaction. in which an amino acid is de-ammonified by anaerobic oxidation yielding also VF A and H2, in conjunction with the reductive de-ammonification of other amino acids consuming the

Table 16.3 Averaged kinetic properties of acidifiers and methanogens

Process Conversion rate gCOD/gVSS.d

Acidogenesis 13 Methanogenesis 3 Overall 2

produced H2. From both reactions NH 3 is released and subsequently acts as a proton acceptor, thus leading to a pH increase. In this reaction there is no net proton production and there is no chance of reactor pH drop.

16.2.1.3 Acetogenesis

The short chain fatty acids (SCFA), other than acetate, which are produced in the acidogenesis step are further converted to acetate, hydrogen gas and carbon dioxide by the acetogenic bacteria. The most imp011ant acetogenic substrates are propionate and butyrate, key-intermediates in the anaerobic digestion process. But also lactate, ethanol, methanol and even H2 and C02 are (homo )acetogenically conve11ed to acetate as shown in Figure 16.3 and Table 16.4. LCF As are converted by specific acetogenic bacteria following the so-called ~oxidation in which acetate moieties are split from the aliphatic chain (Table 16.4). LCF As with uneven C atoms also yield propionate next to acetate. Non-saturated LCFAs like oleate and linoleate are firstly saturated by H2 addition prior to the ~ -oxidation. The acetogenic bacteria are obligate hydrogen producers and their metabolism is inhibited by hydrogen, which immediately follows from the stochiometric conversion reaction , such as for propionate:

3 LIG' = L!Go + RT In [Acetate} [C02 } [ H2 } (l 6.4) [Propionate}

Studies caJTied out on acetogenic conversions have elucidated the required nanow associations between the Hrproducing acetogenic bacteria and the Hrconsuming methanogenic bacteria, thereby regulating the H2 level in their environment. This is of vital importance as these reactions are thermodynamica lly unfavourable, indicated by the positive L':.G0 ' in Table 16.4. From this table it follows that the reactions for ethanol, butyrate,

y K, llm gVSS/gCOD mgCOD/1 lid

0.15 200 2.00 0.03 30 0.12

0.03-0.18 0.12

propionate and the LCF As palmitate will not occur under standard conditions, as the L1G0 ' is positive, and thus the bacterial energy yield is negative.

However, under stabilised digestion conditions the hydrogen partial pressure is maintained at an extremely low level. This can be achieved by an effective uptake of the hydrogen by methanogens or sulphate reducing bacteria. Methanogenic bacteria usually utilize molecular hydrogen in the anaerobic digester so rapidly that the hydrogen partial pressure drops below I o4 atm, which is enough to ensure the actual occurrence of the hydrogen producing acetogenic reaction (Figure 16.6).

This interdependence means that the degradation of higher fatty acids and alcohols largely depends on the activity of electron scavenging organisms such as methanogenic bacteria. Microbial associations in which a Hrproducing organism can grow only in the presence of a H2-consuming organism are called syntrophic

150

~ 0 -100

421

associations. The coupling of formation and use of H2 is called interspecies hydrogen transfer. In a properly functioning methane-producing installation, the partial hydrogen pressure will not exceed I o4 atm and is usually between 104-I0-6 atm. At such a low hydrogen concentration, the degradation of ethanol, butyrate or propionate becomes exergonic and will yield energy for the acetogens.

Similar to the other acetogenic substrates, LCF A conversion is highly endergonic and often limits the entire digestion process (Novak and Carlson, 1970). Trials with upflow anaerobic sludge blanket (UASB) reactors were only partly successful as LCF A tend to absorb to the sludge forming fatty clumps of biomass with little if any methanogenic activity. Expanded bed reactors, in which the LCF A is more evenly distributed over the available biomass were more successful (Rinzema, 1988). Other authors propose in fact to use the absorptive capacity of the sludge and periodically

E ~ propionate + 3 H20 -----+ acetate+ HC03 + 3 H2 + W -,

::::. 0

422

load the s ludge with LCF A after which solid state digestion will convert the absorbed matter to CH4 (Pereira et al., 2004). Such a sequencing bed mode of operation requires multiple reactors to treat a continuous flow wastewater.

16.2.1-4 Methanogenesis

Methanogenic bacteria accomplish the final stage in the overall anaerobic conversion of organic matter to methane and carbon dioxide. During this fourth and last stage of anaerobic degradation of organic matter, a group of methanogenic archea both reduce the carbon dioxide using hydrogen as electron donor and decarboxylate acetate to form CH4 (Figure 16.3). It is only in this stage when the influent COD is converted to a gaseous form that automatically leaves the reactor system. Methanogens are obligate anaerobes, with a very narrow substrate spectrum. Some can only use certain determined substrates such as acetate, methylamines, methanol, formate , and H2/C02 or CO. For engineering purposes , methanogens are classified into two major groups: the acetate converting or aceticlastic methanogens and the hydrogen utilising or hydrogenotrophic methanogens (Table 16.5). Generally, about 70 % of the produced methane originates from acetate as the main precursor. The rest mainly originates from H 2 and C02. The growth rate of the aceticlastic methanogens is very low, resulting in doubling times of several days or even more . The extremely low growth rates explain why anaerobic reactors require a very long start-up time with unadapted seed material and why high sludge concentrations are pursued. Hydrogenotrophic bacteria have a much higher maximum growth rate than the acetoclastic bacteria with doubling times of 4 to 12 hours . Because of this feature and despite the very delicate acetogenic reaction step discussed in the previous section, anaerobic high-rate reactor systems exert a remarkable stability under varying conditions.

Table 16.5 lists two types of aceticlastic methanogens with very different kinetic characteristics.

l j [' )!r

A lso the morphological characteristics of both methanogenic genera are very different as indicated by Figure 16.7.

Figure 16.7 Morphology and appearance of the most important acetotrophic methanogens belonging to the genera Methanosarcina (above) and Methanosaeta (below)

Methanosarcina spp. are characterised by a coccoid shape, appearing in small grape-like clumps, and have a relati ve wide substrate spectrum as they can convert a.o. acetate, H2/C02, methy lamines, methanol , and formate. They have a relatively high llmax and relative low substrate affinity. Methanosaeta spp. are filamentous, appear in large spaghetti like conglomerates can only convert acetate and are kinetically characterised by a low ~Lmax and a very high substrate affinity. Although the llmax of the latter organism is significantly lower, Methanosaeta spp. are the most common acetotrophic methanogens in anaerobic high rate systems based on high solids retention times, such as sludge bed systems and anaerobic filters . The reason for this phenomenon can be attributed to the fact that wastewater treatment

Table 16.5 Most important methanogenic reactions, the corresponding free energy change (CIG 0 ') and some kinetic properties

Functional step

Acetotrophic methanogenesis '

Reaction

Hydrogenotrophic C02 + 4Hl ~ CH-1 + 2Hl 0 methanogenesis

6Go' kJ/mol

-31

-131

~l rnax 1/d

0.12" 0.71 b

2.85

5.8' l.Ob

0.2

' Two different methanogens belonging to 'Methanosarcina spec. and bMethanosaeta spec.

K, Eq. mgCOD/1

30a (16.12) 300b

0.06 (16.13)

\ll

systems always aim at the lowest possible effluent concentrations, while substrate concentrations inside biofilms or sludge granules of the mentioned anaerobic systems approaches ' zero' when bulk liquid concentrations are low. Under such conditions, Methanosaeta spp. species have a clear kinetic advantage over the Methanosarcina spp. (Figure 16.8).

Methanosaeta

Substrate concentration

Figure 16.8 Monod growth curves of the acetotrophic methanogens Methanosarcina spp. and Methanosaeta spp. Both

~lmox and the Monod half saturation constant (K,) of both genera is given in Table 16.5

Once the Methanosaeta spp. dominate the sludge bed, a very effective wastewater treatment system is obtained, reaching extremely low effluent acetate concentrations. Considering the inferior kinetic properties at low substrate concentrations and the inferior adherence properties of Methanosarcina spp., it is advised to keep the effluent acetate concentrations at a very low level during the first start-up of an anaerobic reactor with unadapted seed material.

16.3 PREDICTING THE CH 4 PRODUCTION

The organic pollution can be classified on the basis of solubility (soluble and insoluble organic matter) and/or on the basis of biodegradability.

Both are of great importance for the treatment process. Regarding the enormous variety of organic compounds generally present in wastewater it is impractical and generally also impossible to determine these compounds separately. In order to quantify the organic pollution in practice, use is being made of the fact that these contaminants can be oxidised by strongly oxidizing agents. In wastewater treatment engineering practice two standard tests based on the oxidation of organic material are applied: the biochemical oxygen

demand (BOD) and the chemical oxygen demand (COD) tests (Chapter 3). In both tests , the organic material is oxidised and the amount of oxygen consumed stands for the value of the parameter. In the BOD test it concerns the biochemical amount of oxygen required by the aerobic organisms to oxidize the organic matter. The BOD value therefore is closely related to the biodegradability. For application of anaerobic treatment, it is preferable to use some kind of standardized anaerobic biodegradability test instead of the conventional aerobic BOD test. Tn such an anaerobic test a sample of the wastewater is exposed to an available amount of anaerobic sludge and the total amount of CH4 produced after termination of the digestion process is determined and then related to the amount of organic matter present in the sample. As a certain amount of CH4 is equivalent to a certain amount of COD, we in fact determine the BODanaerobic

Since generall y not all organic pollutants are biodegradable and also part of the organic substrate will be used for cell synthesis, the BOD value generally is substantially lower than the COD value. Latter is particularly the case for the conventional aerobic BOD test, much less for the anaerobic BOD test because of the significantly lower growth yield under anaerobic conditions. Efforts for standardisation are currently being done including ring tests in various laboratories.

In the standardized COD test, which generally uses bichromate as oxidizing medium at an elevated temperature (150C), almost all organic pollutants are completely converted into C02 and H20. On the other hand organic nitrogen present in the contaminants is converted into NH3, whereas organic matter containing quaternary ammonium salts like betaine (trimethyl glycine) stay as well reduced and are ' invisible ' in the COD test.

The total organic carbon (TOC) IS another measurement used, but it is a much less useful parameter. The organic carbon concentration IS measured In the form of carbon dioxide after incineration of the organic material present in a waste water sample . Correction must be made for inorganic carbon, originally present in the sample. The theoretical value of a pure compound follows from Eq. 16.14:

TOC1 = 12n I (1 2n +a+ 1 6b + 14d) (gTOC/gCnHaOb d) (16.14)

424

The COD undoubtedly represents the most important parameter for the concentration of contaminants in wastewater, particularly for industrial wastewaters. This feature in which organ ic matter is almost completely oxidized makes the COD test very suitable for assessment of COD balances. Calculation of the substrate COD and the theoretical quantity of methane produced is presented be low.

The COD of an organic compound CnHaOb can easily be calculated on the basis of the chemical oxidation reaction, assuming a complete oxidation:

I CnHaOb +-;;r 4n + 1- 2b)02 ~ nC02 + (a I 2)H p

(16.15)

Eq. 16.15 shows that 1 "mol" of organic material demands 1/4( 4n+a-2b) moles 0 2 or 8( 4n+a-2b) g02 Hence the theoretical oxygen demand of organic material can be expressed as:

COD, =8(4n +a-2b) / (12n +a +l6b) (gCOD/gCnHaOb) ( 16. 16)

Obviously, with nitrogen containing compounds (proteins and amino acids) Eq. 16. 16 needs to be conected for the number of electrons that will stay with Nand the total weight ofN in the compound.

COD,= 8( 4n +a- 2b -3d) I (1 2n +a+ 16b + 14d) (gCOD/gCnHaObNd) ( 16.17)

From the chemical-oxidation equation for acetic acid,

(16.18)

follows that 1 mol (60 grams) of acetic acid (oxidation number of the C atom is 0) requires 2 moles (64 grams) of oxygen. This means that 1 gram of acetic acid requires 64/60 (1.067) grams of oxygen, consequently 1 gram of acetic acid con esponds to 1.067 gram COD.

The ratio between the COD and TOC values is calculated from :

COD I TOC = 8( 4n +a- 2b- 3d) I ( 12n) = (16.19)

8 I 3+ 2(a-2b - 3d) l ( 3n)

0 t T ""r F le 'I - E lhr , r C ) S r

Table 16.6 summarizes the calculated theoretical values of the COD per unit mass for a number of organic compounds of the type CnHaObNd. The COD per unit mass may be very different for different chemical compounds. In the case of strongly reduced compounds, for example methane, this COD is high, by using Eq. 16.2 for methane (CH4 i.e. n=l , a=4, b=O, d=O) in Eq. 16.4 one calculates:

CODCH, = 8( 4 1 + 4-2 0-3 0) I (16 .20)

(121 +4 +16 0+14 0) = 4gCOD I gCH4

It is clear that the ratio of COD and TOC differs substantially for the various compounds. This is explained by the differences in the average oxidation state of the organic carbon. The carbon oxidation state (C-ox. state) of carbon can va1y from -4 (the most reduced state of carbon, as found in CH4) to +4, the most oxidized as found in C02 Figure 16.9 depicts for a number of compounds their mean C-ox. state in relation to the theoretical composition of the biogas produced which obviously yields a linear conelation (Table 16.6).

~ c::

~

100 CH4 0 COz

-~ SO CH4 ~ COz c::

E ~ E 0 u

Methanol

Carbohydrates, Acetic acid

Citric acid

Formic acid Carbon monoxide

Oxalic acid

0 CH4 +------,------.,-------'?-- Urea 100 COz _

4 2 4

Mean oxidation state of carbon

Figure 16.9 Theoretical composition of the biogas produced in rela tion t o the mean oxidation state of the carbon in specific substrates, assuming complete mineralization of the substrate

The lower the average oxidation state of the carbon m a compound (i.e . the more negative), the more oxygen can be bound by the compound, and consequently the higher is its COD value.

Since organic-N is converted into NHrN in the COD test (consequently the amount of N in the compound should be accounted for as being in its reduced form, i. e. having taken up 3 electrons) , and one atom of H

. b T c t 425

Table 16.6 Stoichiometric values of COD and TOC per unit mass for different pure organic compounds C, H,ObN d, the COD/TOC values and the mean carbon oxidation state for these compounds and the estimated CH 4 % in the biogas

Compound

Methane Ethane Methanol Ethanol Cyclohexane Ethylene Palmitic acid Acetone Ethylene glycol Benzene Betaine Glycerine Phenol Lysine Phenyl alanine Insuline Glucose Lacitc acid Acetic ac id Citric acid Glycine Formic acid Oxalic acid Carbondioxide

n a b d

I 4 0 0 2 6 0 0

4 0 2 6 1 0 6 12 0 0 2 4 0 0 16 32 2 0 3 6 0 2 6 2 0 6 6 0 0 5 11 2 I 3 8 3 0 6 6 1 0 6 14 2 2 9 11 2

254 377 75 65 6 12 6 0 3 6 3 0 2 6 2

2

4 8 5 2 2 0

2 7 2 2 4 2

0 0 I 0 0 0

4 3.73

1.5 2.09 3.43 3.43 3.43 2.21 1.29 3.08 1.64a 1.22 2.38 1.53 1.94 1.45 1.07 1.07 1.07 0.75 0.64 0.35 0.18

0

gTOC/ COD/ TOC g CnHaObNd

0.75 5.33 0.8 4.67

0.38 4 0.52 4 0.86 4 0.86 4 0. 75 3.83 0.62 3.56 0.39 3.33 0.92 3.33 0.5 1 3.2 0.39 3.11 0.77 3.11 0.49 3.11 0.65 2.96 0.53 2. 72

0.4 2.67 0.4 2.67 0.4

0.38 0.32 0.26 0.27 0.27

2.67 2 2

1.33 0.67

0

C-ox. state

-4 -3 -2 -2 -2 -2

-1.75 -1.33

-1 -I

-0.8 -0 .67 -0 .67 -0.67 -0.44 -0.08

0 0 0

1 2 3 4

100 87 .5

75 75 75 75 72 67

62.5 62.5

60 58 58 58 56 51 so so so

37.5 37.5

25 12.5

0 a Calcu lated COD. Theoretical: with standardi sed bi-chromate COD test no COD will be measured

provides one electron and one atom of 0 will take up two electrons, the average oxidation number of the C atom in a compound CnHaObNd follows from:

C- ox.state = (2b- a +3d) I n (16.21)

The number of electrons made free per atom C in the complete oxidation of CnHaObNd amounts to:

4- ( 2b +3d- a) I 11 = 4 + (a- 2b- 3d) I 11 (16.21)

Consequently the number of molecules 0 2 required for the oxidation amounts to:

n + 1 I 4a - 1 I 2b- 3 I 4d (16.22)

Therefore the equation for complete chemical oxidation of this compound is: C11 H 0 0 6Nd +(11+a I 4 -b I 2-3d I 4)02 ~ ( 16.23) nC02 + (a I 2 -3d I 2 )H20+ dNH3

In case the compound (CnHaObNd) is completely biodegradable and would be entirely converted by the anaerobic organisms (no sludge y ield) into CH4, C02 and NH3, the theoretical amount of methane gas (and C02) produced can be calculated using the Buswell equation:

C11 H 0 0 6Nd +(n-a l 4-b l 2+3d I 4)H20~ ( n l 2+a l 8-b l 4-3d i 8)CH4 + ( 16.24) (n I 2-a I 8+b I 4 +3d I 8)C02 +dNH3

The COD provides the correct infonnation concerning the oxidation state of the compound, consequently the amount of methane that can be produced from it (Table 16.6, Figure 16.9). The only exceptions are th e quaternary ammonium salts such as the a lready mentioned betaine, wh ich stays reduced in the laboratory COD test. Therefore, the COD is generally accepted as the most adequate parameter to quantifY the concentration of organ 1c materi a l and

426

certainly not the TOC. For predicting the relative amount of CH4 in the produced biogas when the exact composition of the organic matter is unknown, the COD/TOC ratio is a very useful tool. The latter is based on the linear conelation between the mean oxidation state and the COD/TOC ratio (Figure 16.1 0).

120 +2 0 2 4

100 'C' mean oxidation state CH,

Ethane

~ 80 ~ "" ~ 60 Glycerine, Ph enol c Acetate , Glucose , Lactic acid Cl u 40

Ci tric acid , Glycine

20 Formic acid

COD/TOC ratio

Figure 16.10 Expected CH 4 % in the produced biogas as a function of the COD/TOC ratio: CH 4% = 18.75 COD/TOC

ln the presence of specific inorganic electron acceptors like nitrate , sulphate or sulphite, the production of methane will decrease, due to the occurrence of a.o. the following reactions:

( 16.25)

( 16.26)

For wastewaters containing an excess of organ ic electron acceptors with respect to the amount of nitrate (N03-) , nitrite (N02-) , su lphate (SO/ ) or sulphite (SO/ ) present, a complete removal of these electron acceptors (oxygen donors) may occur. Since the solubility of H2S in water considerabl y exceeds that of CH~, a substanti al lower COD removal from the water phase will be obtained in case the ,,aste\\ater contains sulphate.

The quantity of C02 present in the biogas produced generally is significantly lower than follows from the Buswell equation or the COD/TOC ratio as depicted in Figure 16.1 0. This is because of (a) the relatively high solubility of C02 in water and (b) because part of the C0 2 may become chemically bound in the water phase due to the formation of ammoni a in the anaerobic conversion of nitrogen containing organic compounds and cations which were present in the wastewater as salts ofVFA, SO/ , N0 3-.

, , r

IMPACTS OF ALTERNATIVE ELECTRON ACCEPTORS

Bacterial conversions under anoxic conditions

Anaerobic digesters contain mixed microbial communities. Besides the methanogenic association described before, other bacteria are present which can compete with the methanogens for methanogenic substrates (Table 16. 7). The listed bacteria have different microbial respiration systems and can use different electron acceptors such as oxygen (02) by (facultative) aerobic bacteria, nitrate (N03-) by denitrifiers, sulphate (SO/ ) or sulphite (SO/) by sulphate reducing bacteria and iron (Fe3+) by iron reducers. Anoxic means that oxygen in the form of oxygen gas (02) is not available as an electron acceptor.

16-4.1.1 Sulphate reduction ln the presence of sulphate, sulphite or thiosulphate, sulphate reducing bacteria (SRB), which have a much wider substrate spectrum, are ab le to use several intermediates of the anaerobic mineralisation process (Table 16. 7). These bacteria convert sulphate into hydrogen sulphide. Besides the direct methanogenic substrates such as molecular hydrogen (H2) , fom1ate, acetate, methanol and pyruvate, SRB can also use propionate, butyrate, higher and branched fatty acids, lactate, ethanol and higher alcohols, fumarate, succinate, malate and aromatic compounds (Colleran et al., 1995). Hence, the main inte1mediary products of the anaerobic degradation process (H2/CH3COO-) can be converted by both SRB, methanogens and/or obligate hydrogen producing bacteria (OHPB). Because these three groups of bacteria operate under the same environmental conditions (pH, temperature), they will compete for the same substrates . The outcome of this competition depends on the conversion kinetics (see Section 16. I 0).

lf organic material is oxidised via sulphate reduction, 8 electrons can be accepted per molecule of sulphate. Since one molecule of oxygen can only accept 4 electrons, the electron accepting capacity of 2 moles of 0 2 equals I mol of SO/ , equivalent to 0.67 g of 0 2 per g SO/ . This means that for waste streams with a COD/sulphate ratio of 0.67 , there is theoreticall y enough sulphate available to completely remove the organic matter (COD) via sulphate reduction. For COD/sulphate ratios lower than 0.67, the amount of organic matter is insufficient for a complete reduction of the sulphate present and extra substrate then should be

b 427

Table 16.7 Stoichiometry and change of free energy I':.G"' (kJ /mol substrate) of hyd rogen and acetate conversion under different cond itions

Reaction Aerobes H2 + 0.5 Or -> H20 CH3coo + 2 o2 ___. 2 Hco3- +IT Den i tri fi ers H2 + 0.4 N03- + 0.411---> 0.2 N2 + 1.2 H20 CH3Coo + 1.6 NO;-+ 0.6 H+---> 2 HCO;- + 0.8 N2 + 0.8 H20 Fe3+ reducing bacteria H2 + 2 Fe3- --->2 Fe2- + 211 CH3Coo + 4 Fe

3+ + 4 H20---> 4 Fi + + 5 F + 2 H C03 Sulphate reducing bacteria H2 + 0.25 SO/ + 0.25 W---> 0.25 HS + H20 CH3Coo +SO/---> HS + 2 HCOJ-Methanogens H1 + 0.25 HC03- + 0.25 /--{"---> 0.25 CH-1 + 0. 75 H20 CH3COO- + H20---> CH4 + HCO;-

added if removal of sulphate is the objective of the treatment. On the contrary, for wastewaters with a COD/sulphate ratio exceeding 0.67 , a complete removal of the organic matter can only be achieved if, in addition to sulphate reduction, methanogenesis also occurs.

In the presence of sulphate, organic matter is not necessarily degraded less eas il y, but compared to methane, hydrogen-sulphide has the great disadvantage that it disso lves much better in water than methane. This means that, for the same degree of organi c waste degradation , a lower quantity of COD will be reduced in wastewater containing sulphate. Su lphide production can further cause the following process technical problems during anaerob ic di gestion:

H2S is toxic to methanogenic bacteria (MB), acetogenic bacteria (AB) and SRB. In case of methanogenic treatment of the waste-stream, some of the organic compounds in the wastewater will be used by SRB rather than MB and are therefore not converted into methane. This results in a lower methane yield per unit of degraded organic waste and, therefore, negatively affects the overall energy balance of the process. Moreover, the quality of the biogas is reduced since a part of the produced sulphide ends up as H2S in the biogas. Removal of H2S from the biogas is therefore usually required.

The produced sulphide has a bad smell and can cause corrosion problems to pipes, engines and boilers. Thus, the maintenance costs of the installat ion

L'.G 0 ' (kJ/mol substrate) Eq.

-237 (16 .27) -844 (16 .28)

-224 (16.29) -792 ( 16.30)

-228 (16.31 ) -352 (16.32)

-9.5 (16.33) -48 ( 16.34)

-8.5 (16.35) -31 (16 .36)

increase and extra investment costs are necessary to avoid these problems.

Part of the sulphide will be present in the effluent of the anaerobic reactor. As mentioned above, thi s results in a lower overall treatment efficiency of the anaerobic reactor system, as sulphide contrib utes to the wastewater COD (per mole of sulphide two moles of oxygen are required for a complete oxidation into sulphate). Moreover, sulphide can upset the treatment efficiency of the aerobic post treatment system, e.g. algal blooming in lagoons or activated sludge bulking. Thus, an extra post treatment system to remove the sulphide from the wastewater may be required.

Based on their substrate consumption, SRB may be classified into the following three groups:

1) hydrogen ox idising SRB (HSRB) 2) acetic acid oxidising SRB (ASRB) 3) fatty acids oxodising SRB (FASRB)

In the last group, two oxidation patterns can be di stinguished:

CH 3CH 2COOH + 2H;O~ CH3COOH +3H2 +C02 (0HPB )

CH3CH 2COOH +0.75SO/- ~ CH3COOH +0.75s

2- +C02 ( FASRB )

(16 .37)

(16.38)

428

CH3CH2COOH + 1.75SO/- --+

175S 2- +3C02 +3H;O(FASRB) ( 16.39)

Some SRB are capable of completely oxidising VF A to C02 and sulphide as end products. Other SRB lack the tricarboxylic acid cycle and catTy out an incomplete oxidation of VF A with acetate and sulphide as end-products. In the latter case, acetic acid is excreted in the medium. It should be further noticed that incomplete oxidation of propionic acid by an SRB yields the same degradation products as the conversion by the OHPB and HSRB. Hence, it is not possible to deduce from mass balances which bacteria carry out his con ersion. C~ll',- o!,\1.("'t\\l '-\yt'-..,O~fvt. r~" vL~'i\ L\t:'fiA HS rz \!:,- "'IA'Io{\i'-t bx'6."s'"' \ '5 11'\1 1-z.. 'I( b~-I.JMI(.JL.U [lv-./lt'{ h~ ,

Tf SO/ is present in the wastewater, SO/ reduction by SRB cannot be prevented. Several attempts were made to try to steer the competition in a single reactor system but were unsuccessful. On the other hand, several technological solutions are available on the market that are directed to lower the H2S concentration in the anaerobic reactor to minimise the toxicity of the MB (Figure 16.11).

16-4.1 .2 Denitrification

In general, no denitrification occurs during anaerobic purification and digestion. Organically bound nitrogen will be converted into ammonium. Denitrification can only be expected if the influent contains nitrate (see Chapter 5).

Denitrification is mediated by denitrifying micro-organisms, i.e. chemoheterotrophic bacteria which are capable of oxidising organic matter with nitrate . Nitrate is then converted via nitrite and nitrogen oxide into N2 gas. Generally, denitrifying micro-organisms prefer oxygen as an electron acceptor, as the latter compound yields more energy (Table 16. 7). In aerobic purification processes, they start to use nitrate as soon as 0 2 is depleted to cope with the organic load. ln an activated-sludge plant, denitrification. will normally occur only at a dissol ved 0 2 concentration of 1 mg/1 or below.

Denitrification is a heterotrophic process requiring an electron donor. The stoichiometry of methanol oxidation with nitrate and nitrite occurs according to the following reaction equation:

( 16.42)

5CHpH + 6N03---+ 3N2 +4HC03- +CO/ - +8H20 (16.43)

These reaction equations show that denitrification will result in a pH increase (carbonate production).

A

Wastewater

B

Biogas

Methanogenesis Sulfate reduction

t

Gas recycle

Effluent Gas scrubbing

l Effluent recycle Wastewater

Methanogenesis Sulfate reduction 1-----1 Sulfide removal I 10 Effluent

'----------' E Stn ppmg AbsorptiOn

810leog1Cal ox1dat10n to so c

Wastewater Acidification Sulfate reduction ~~Effluent

Figure 16.11 Technological solutions to decrease the H,S concentration in the anaerobic reactor. (A) enhanced H,S stripping by biogas recycling and sulphide stripping in the gas line, (B) H,S removal in a (micro)aerobic post treatment system and recirculation of the treated effluent to the anaerobic reactor influent for dilution, (C) combined pre-acidification and sulphate reduction with sulphide removal step for lowering the S content in the anaerobic reactor. In the latter approach most of the H,S will be stripped in the acidification step owing to the low prevailing pH

WORKING WITH THE COD BALANCE

Like any biological system an anaerobic treatment process must be monitored for relevant parameters, and measurements must be evaluated for adequate operation and control. Section 16.3 discusses the usefulness of the COD as the control parameter for anaerobic systems. The reason for this is that in contrast to aerobic systems there is no COD destn1ction in an anaerobic reactor. During anaerobic treatment the COD is only 're-arranged'. Complex organic compounds are broken down in more simple intermediates and eventually mineralised to CH4 and C02. All COD that entered the system ends up in the end-product CH4 , minus the COD that is incorporated in the new bacterial mass. Since a perfect mass balance can be made by only using the COD as a parameter, the COD is therefore generally taken as a control tool to operate an anaerobic system:

COD,/1 = CODOll( (16.44)

For practical purposes Eq. 16.44 should be expanded to the various outlets of the anaerobic reactor as depicted in Figure 16.12.

For identifying the fate of COD in an anaerobic reactor detailed analyses of the gaseous, liquid and solid outlets should be performed (Table 16.8).

CODgas

f CODinfluent

..... ~ CODetfl uent

1 CODstudge

CODtnfluent = CODeffluent + CODgas + CODsludge

Figure 16.12 COD balance of an anaerobic reactor. By differentiating the COD fractions of gas, liquid and solids, the missing parameters can be estimated from the more easily measurable parameters

Based on the basic influent characteristics, i.e. flow rate and COD concentrations, and information on the biodegradabili ty of the COD, the expected CH4 production rate can be easi ly estimated. From section 16.3.1. we can derive that:

( 16.45)

which means that 22.4 m3 CH4 (STP) requires 2 moles of 0 2 (COD), which equals 64 kg COD. Therefore, theoretically, 1 kg COD can be converted in 0.35 m3 CH4.

430

Similarly, the theoretical COD equivalent for l kg ' bacterial VSS', with an estimated composition of C5H 70 2N, can be calculated as 1.42 kgCOD/kgVSS. Having both the final products CH4 and newly grown bacteria expressed as COD, the balance can be made if influent and effluent are properly measured.

Often ' gaps ' in the COD balance occur which can be attributed mostly to the ' loss of electrons' when these are channelled to ox idi sed an ions like SO/ and N03-, as explained in section 16.4. Therefore, in this case, for clos ing the COD balance either a ll reduced gases should be taken into account or the concentration of electron acceptors needs to be measured. Tt should be realised that soluble COD containing gases like H2S, will be present in the effluent. In this example, organic COD is converted into inorganic COD of which a pH dependent frac tion will end in the biogas while the remainder wi ll stay in the effluent.

An other frequently cited cause for a COD gap is the entrapment or accumulation of COD in the sludge bed, sometimes drastically changing the stochiometric value of 1.42 kgCOD/kgVSS. The latter is particularly true during the treatment of fat- or LCF A-containing wastewater. With these substrates, COD removal efficiencies are genera ll y very high, but low CH-1 production rates lead to huge gaps in the balance. In thi s example, the COD gap indicates severe long-term operational problems. The accumulating solids will deteriorate the SMA of the s ludge, finally resulting in a complete failing of the anaerobic process.

Operating an anaerobic reactor using the COD balance as a tool to monitor reactor performance gives the operator vital information about the functioning of

< T 'rl li -,

the system. Adequate action can be undertaken before iiTeversible deterioration occurs. Also, the impact of altemative electron acceptors on the CH4 production rate can be easily assessed while based on the gas production and effluent COD values, an estimate can be made of the amount of newly grown and entrapped biomass.

16.6 IMMOBILISATION AND SLUDGE GRANULATION

The key for modem high-rate biotechnology, whatever systems will be considered, is the immobilization of proper bacteria. The required high sludge retention in anaerobic treatment systems is based on immobilization, though it is not just a matter of immobilizing bacteria but of well balanced bacterial consortia. Regarding the occurrence of various syntrophi c conversion reactions in the anaerobic conversion of most organic compounds, the detrimental effect of higher concentrations of spec ific intermediates and the strong effect of environmental factors like pH and redox potential, the development of balanced bacterial consortia is a pre-requisite for a proper anaerobic treatment system. Significant progress in the knowledge of the fundamentals of the immobilisation process has been made smce the development and successful implementation of high rate anaerobic treatment systems in the seventies . Immobili sation may occur on inert support material mounted in a fi xed matrix in so-called anaerobic filters (AF) , which are operated both in upflow and in downflow mode . The matrix can also be free floating like in moving bed bioreactors and fluidi zed bed (FB) systems. Tf no inert support material is used, a so-called 'auto-immobilisation' will occur, which is understood as the immobili sation of bacteria on

Table 16.8 Various COD fractions and their fate in an anaerobic reactor system. Number of dots indicates the relative importance of the indicated COD fraction in the respective compartment ( influent, effluent, sludge, biogas)

COD fraction Influent Effluent Sludge Gas Soluble organic Soluble inorganic Suspended organic Suspended inorganic Colloidal Absorbed

Entrapped CH -1 H2 HzS N2 Newly grown biomass

Ar' E' b

themselves in bacterial conglomerates, or on very fine inert or organic particles present in the wastewater. The bacterial conglomerates will mature in due time and form round shape granular sludge _

With respect to immobilization, particularly the phenomenon of granulation has puzzled many researchers from ve1y different disciplines_ Granulation in fact is a completely natural process. It will proceed in all systems where the basic conditions for its occurrence are met, i_e, on mainly soluble substrates and in reactors operated in an up-flow manner with hydraulic retention times (HRT) lower than the bacterial doubling times. Owing to the very low growth rate of the cmcial aceticlastic MB, particularly under sub-optimal conditions, the latter conditions are easily met. Sludge granulation also was found to occur in reversed flow Dorr Oliver Clarigesters applied in South Africa since the fifties of the last centl1ly However, this only became apparent by observation of sludge samples taken from such a digester in 1979. Surprisingly enough no attention was given to the characteristics of the Clarigester sludge such as size, form and the mechanical strength, density and porosity of sludge floes/aggregates. Despite all the efforts made to develop systems with a high sludge retention nobody apparently noticed that major part of the sludge consisted of a granular type of sludge. While studying the start-up and feasibility of anaerobic upflow filters, Young and McCarty (1969) already recognized the ability of anaerobic sludge to form very well settleable aggregates. These granules were as large as 3.1 mm in diameter and settle readily. In AF experiments with potato starch wastewater and methanol solutions conducted in the Netherlands similar observations were made (Lettinga et al. , 1972, 1979). Whereas the interest in AnWT in USA and South Africa diminished, large emphasis on developing industrial scale systems was put in the Netherlands, where the instalment of new surface water protection acts coincided with the world energy crises of the seventies_ As a result, increasing emphasis could be afforded on applied and fundamental research in this field , particularly also on the phenomenon of sludge granulation_ A worldwide growing interest occurred from both the engineering and the microbiological field. As a result , the insight in the mechanism of the sludge granulation process for anaerobic treatment has been elucidated suffici ently , at least for practical application (e.g. de Zeeuw. 1982; 1987; Hulshoff Pol and Lettinga, 1986; Wiegant and de Man, 1986; Beeftink and Staugard, 1986; Hulshoff Pol et al. , 1987, 2004; Wu, 1987; Dolfing, 1987; Wu er al.,

431

1991; Grotenhuis, 1992; van Lier et a{ , 1994; Thaveesri et al. , 1994; Fang et ar, 1994). Granulation can proceed under mesophilic, thermophilic and psychrophilic conditions. It is of huge practical importance to Improve the insight in fundamental questions concerning the growth of mixed balanced cultures. This will lead very likely to the application of the process for the degradation of a large variety of (difficult) chemical compounds. These challenging questions need to be attacked jointly through the efforts of process scientists and microbiologists.

16.6.1 Mechanism underlying sludge granulation In essence, sludge granulation finds its ground in the fact that bacterial retention is imperative when dilution rates exceed the bacterial growth rates. Immobilization further requires the presence of support material and/or specific growth nuclei. The occurrence of granulation can be explained as follows:

1) Proper growth nuclei, i.e. inert organic and inorganic bacterial carrier materials as well as bacterial aggregates, are already present in the seed sludge.

2) Finely dispersed matter, including viable bacterial matter, will become decreasingly retained, once the superficial liquid and gas velocities increase, applying dilution rates higher than the bacterial growth rates under the prevailing environmental conditions . As a result film and/or aggregate formation automatically occurs_

3) The size of the aggregates and/or biofilm thickness are limited, viz. it depends on the intrinsic strength (binding forces and the degree of bacterial intertwinement) and the external forces exerted on the particles/films (shear stress). Therefore at due time, particles/films will fall apart, evolving the next generation_ The first generation(s) of aggregates, indicated by Hulshoff Pol et al. (1983) as 'filamentous' granules mainly consist of long multi-cellular rod shaped bacteria. They are quite voluminous and in fact more flock than granule_

4) Retained secondary growth nuclei will grow in size again, but also in bacterial density. Growth is not restricted to the outskirts, but also proceeds inside the aggregates . At due time they will fall apart again, evolving a third generation, etc.

5) The granules will gradually 'age' or 'mature'. As a result of this process of maturing the voluminous 'filamentous granules ', predominating during the initial stages of the granulation process, will disappear and become displaced by dense 'rod'

432

granules. In a matured granular sludge, filamentous granules generally will be absent.

During the above described selection process, both organic and hydraulic loading rates gradually increase, increasing the shear stress inside the system. The latter results in firm and stable sludge aggregates with a high density and a high superficial velocity. Figure 16.13 pictures the course in time of the in-reactor sludge concentrations, expressed as gVSSi l, and the applicable organic loading rate. The start is accomplished when the design loading rate is reached. For mainly soluble wastewaters which are partly acidified, granular sludge will be easily cultivated.

Table 16.9 lists some common characteristics of methanogenic granular sludge.

With respect to the granulation process, essentially there do not exist any principle differences between a UASB reactor, seeded with digested sewage sludge, and an upflow reactor with inert free floating support material like the FB reactor, which uses sand particles or pumice as camer material for the in-growing biomass. Granulation indeed can proceed quite well in a FB system, provided the reactor is operated with a moderate shear on the particles, i.e ., in such a mode that biofilms can grow sufficiently in thickness and/or different

d

particles can grow together. Full scale experiences have shown that complete fluidization is not required and is in fact is detrimental in achieving stable and sufficiently thick biofilms. At present the expanded granular sludge bed (EGSB) reactors are of much interest for commercial applications than the more expensive FB systems (see also Section 16.7.2.4).

= 15 ~ > ~

phase l phase 11

r __ J _______ _ ~---0 ___ _I ___ J

0 10 20 30 40

phase Il l

r------------r-----

___ [ ________ _

0 50 60 70 80 90 100

Time (d) Figure 16.13 Sludge dynamics during the first start-up of a UASB reactor. Phase 1: Applied loading rate

16.7 ANAEROBIC REACTOR SYSTEMS

Anaerobic reactors are in use since the 19'" century, when Mouras and Cameron developed the automatic scavenger and the septic tank to reduce the amounts of solids in the sewerage system.Aithough at a very poor rate, the first anaerobic stabilisation processes occurred in the tanks that were designed for intercepting the black-water solids. The first anaerobic reactor was developed in 1905 when Kart Imhoff designed the Imhoff tank, in which solids sediments are stabilised in a single tank. The actual controlled digestion of entrapped solids in a separate reactor was developed by the Ruhrverband, Essen-Relinghausen in Germany.

In the same decades, Buswell stmt ed to adopt the same technology for treating liquid wastes and industrial wastewater. All these systems can be characterised as low rate systems since no special features were included in the design to augment the anaerobic catabolic capacity. The process feasibility of these systems was very much dependent on the growth rate of the anaerobic consortia. As a result, reactors were very big and very fragi le in operation. In the final decades of the 19'h century also some first trials of upward flow fixed film reactors were performed, but it was too early to make these systems successful (McCarty, 200 I). Also the anaerobic pond can be regarded as a low loaded anaerobic treatment system. Anaerobic ponds are often constructed in conjunction with facultative and maturation ponds. The applied loading rate to anaerobic ponds ranges between 0.025-0.5 kgCOD/m3 d, while using pond depths of 4 m. The big disadvantages of anaerobic ponds are problems related to odour as these systems easily become overloaded. Also the loss of energy rich CH4 to the atmosphere is a recognised di sadvantage.

16.?.1 High-rate anaerobic systems

One of the major successes in the development of anaerobic wastewater treatment was the introduction of high-rate reactors in which biomass retention and liquid retention are uncoupled. Contrary to aerobic processes, in an anaerobic or anoxic (denitrification) process, the maximum permissible load is not goYemed by the maximum rate at which a necessary reactant can be supplied (e.g. oxygen during aerobic processes). but by the amount of viable anaerobic biocatalysts or the anaerobic bacteria which are in full contact with the wastewater constituents. In anaerobic high- rate systems, high sludge concentrations are obtained by physical retention and or immobilisation of anaerobic

433

sludge. High biomass concentrations enable the application of high COD loading rates, while maintaining long SRTs at relatively short HRTs. Different high- rate systems were developed over the last three decades including the anaerobic contact process (ACP), anaerobic filters, the UASB, FB and EGSB reactors and the baffled reactors.

To enable an anaerobic reactor system to accommodate high organic loading rates for treating a specific wastewater, the following conditions should be met:

High retention of viable sludge in the reactor under operational conditions. The higher the amount of sludge retained, the higher will be the loading potential of the system. Therefore, it is necessary to cultivate a well settleable or immobilized biomass, and that the sludge will not deteriorate in this respect.

Sufficient contact between viable bacterial biomass and waste water. In the case where part of the sludge retained in the reactor remains deprived of substrate, this sludge is of little if any value.

High reaction rates and absence of serious transport limitations. It is c lear that the kinetics of the degradation processes are a factor of great importance. It is essentia l that metabolic end-products can easily escape from the aggregate. The size of the biofilms should remain relative ly small and the accessibility of the organisms inside the biofilm should be high.

The viable biomass should be sufficiently adapted and/or acclimatized. For any wastewater subjected to treatment, the sludge should be enabled to adapt to the specific characteristics of the concerning wastewater.

Prevalence of favou rable environmental conditions for all required organisms inside the reactor under all imposed operational conditions, focusing on the rate limiting steps. It should be emphasi zed here that this condition doesn't mean that the circumstances should be similar at any location within the reactor and at any instant. As a matter of fac t even the contrary is uue. Regarding the fact that a large variety of different organisms are imo!Yed in the degradation of more complex compounds, the e:-;istence of micro-niches \\ithin the system is an abso lute pre-requisite. Only in this way can the required flourishing grO\nh of the required very different organisms be achie\'ed. It should be noticed that pmticularly in the interior of biofilms and

434

Completely mixed Physical retention Immobilised biomass Enhanced contact

Biogas Biogas Biogas Biogas t t t ~ CD Cffl""_~ _ _ ln._fl"'"~ cp-J"'"'

Relative capacity: 1 Relative capacity: 5 Relative capacity: 25

Influent t Relative capacity: 75

Figure 16.14 Relative loading capacity of different AnWT systems. Maximum applied loading rates under full scale conditions reach about 45 kgCOD/m3.d applying enhanced contact in EGSB type systems

granules, the concentration of substrates and metabolites are low enough to allow even the very endergonic acetogenic reactions to proceed, e.g. the oxidation of propionate at the very low hydrogen concentrations.

As mentioned above, Stander in South Africa and Schroepfer and coworkers were amongst the first to recognize the impot1ance of maintaining a large population of viable bacteria in the methanogenic reactor. On the other hand the idea certainly was not completely new at that time, because the need of the presence of a high viable biomass concentration already was applied in full scale aerobic treatment systems in use in the early fifties and before. It therefore could be expected that supporters of the 'anaerobic concept' would try out the 'aerobic activated sludge' concept for anaerobic wastewater treatment. The anaerobic contact process by Schroepfer et al. ( 1955) indeed tumed out to be reasonably successful for the treatment of higher strength industrial wastewaters. With a few exceptions, hardly any at that time would think that anaerobic treatment ever could become feasible for low strength wastewaters. Regarding the problems experienced with

the various versions of the anaerobic contact process, only very few even believed anaerobic treatment could become applicable for treating medium strength wastewater. However in the sixties and seventies the situation changed rapidly, and in the nineties the anaerobic treatment concept even was shown feasible for very low strength wastewaters at low ambient temperatures. These unforeseen developments can be attributed to superior methods of sludge retention, based on sludge immobilization. Figure 16.14 illustrates the development of high rate reactor systems and the impact of improved sludge retention and enhanced contact on the applicable organic loading rates. While the first trials of Buswell did not reach loading rates of 1 kgCOD/m3 d, modern AnWT systems are sold on the market with guaranteed loading rates exceeding 40 kgCOD/m3.d.

At present, most applications of AnWT can be found as end-of-the-pipe treatment technology for food processing wastewaters and agro-industrial wastewater. Table 16.10 lists the various industrial sectors where the surveyed 2,266 reactors are installed. It should be noticed that the number of anaerobic applications in the

Table 16.10 Application of anaerobic technology to industrial wastewater. Total number of registered worldwide installed reactors = 2,266, census January 2007, after van Lier (2007) (see also Figure 16.2) Industrial sector Type of wastewater Nr. of reactors % Agro-food industry Sugar, potato, starch , yeast, pectin, citric acid, cannery, 816 36

confectionary, fruit , vegetables, dairy, bakery Beverage Beer, ma\ting, soft drinks, wine, fruit juices, coffee 657 29 Alcohol distillery Can juice, cane molasses, beet molasses, grape wine, grain, 227 10 fruit Pulp and paper industry Recycle paper, mechanical pulp, SSC, sulphite pulp, straw, 249 \I bagasse Miscellaneous Chemical , pharmaceutical , sludge liquor, landfill leachate, 317 14

acid mine water, municipal sewage

non-food sector is rapidly growing. Common examples are the paper mills and the chemical wastewaters, such as those containing fom1aldehyde, benzaldehydes, terephthalates, etc. (Razo-Flores et al. , 2006). The latter is surprising since it is pa11icularly difficult for the chemical industries to enter with anaerobic technology, owing to the general prejudices against biological treatment and anaerobic treatment in particular. With regard to the chemical compounds it is of interest to mention that certain compounds, such as poly ch loro-aromatics and poly nitro-aromatics as well as the azo-dye linkages can only be degraded when a reducing (anaerobic) step is introduced in the treatment line. Anaerobics are then complementary to aerobics for achieving full treatment.

Only very recently, high-rate AnWT systems were developed for treating cold and very low strength wastewaters . T n addition to municipal sewage, many industrial wastewaters are discharged at low temperatures, e.g. beer and maltery wastewaters. Full scale results so far show that any of the cited wastewaters are anaerobically treated using common seed materials, illustrating the robustness and flexibility of the anaerobic process.

16.7.2 Single stage anaerobic reactors 16.7.2.1 The Anaerobic Contact Process (ACP) As explained in section 16. 7.1 , processes employing external settlers and sludge return are known as the anaerobic contact process (ACP), see Figure 16.15.

Recycle sludge

Excess sludge

Effluent -----.

Figure 16.15 Anaerobic contact process, equipped with flocculator or a degasifier unit to enhance sludge sedimentation in the secondary clarifier

The various versions of the first generation of 'high rate' anaerobic treatment systems for medium strength wastewaters were not very successful. In practice, the main difficulty appeared to be the separation of the sludge from the treated water. These difficulties can be mainly due to the fact that a too intensive agitation in the bio-reactor was considered necessary. The idea was that the more intensive the mixing, the better would become the contact between sludge and wastewater.

435

However, in that time no consideration was given to the quite detrimental effect of intensive mixing on the sludge structures, viz. its settleability and the negative impact on the presence of balanced micro-ecosystems, i.e. , syntrophic associations (Section 16.2.1.3).

Various methods for sludge separation have been tested and/or employed in the different versions of the ACP. These methods include vacuum degasification in conjunction with sed imentation, the addition of organic polymers and inorganic flocculants, centrifugation and even aeration (in order to stop digestion). However, the results were usually unsatisfactory. At present, with the current knowledge on anaerobic digestion technologies, a more gentle and intermittent mode of mixing is applied. With such an approach, the sludge will acquire and keep excellent sedimentation properties, and the anaerobic contact process can certain ly make a valuable contribution to environmental protection and energy recovery, particularly with wastewaters containing high fractions of suspended solids and semi liquid wastes. If well designed, modem ACP may reach organic loading rates of l 0 kgCOD/m3 .d.

16.7.2.2 Anaerobic Filters (AF) The modem version of uptlow anaerobic filter (UAF) was developed in the USA by Young and McCarty ( !964, 1982) in the late sixties. The sludge retention of the UAF is based on:

the attachment of a biofilm to the solid (stationary) carrier material,

the sedimentation and entrapment of sludge pm1icles between the interstices of the packing material, formation of very well settling sludge aggregates.

Initially, a suitable carrier material for the systems was hard to find (Young, 1991 ). Various types of synthetic packing have been investigated and natural materials such as gravel, coke and bamboo segments as well. It turned out that the shape, size and weight of the packing material are impm1ant aspects. Also the surface characteristics with respect to bacterial attachment is impo11ant. Moreover, it was found that the bed should remain open of structure, viz. providing a large void fraction. Applying proper support material AF systems are rapidly started, owing to the efficient adherence of anaerobic organisms to the inert carrier. The ease of starting up the system was the main reason for its popularity in the eighties and nineties. Problems with UAF systems in particular generally occur during long-term operation. The major disadvantage of the UAF

UASB 50%

19812007, N = 2266

IC 15%

I \ LAG 4%

Reference HYBR 3% incomplete 1% FB 2%

Expanded Bed Reactors: 29% Granular Sludge Bed Based: 77%

B l O)i' c ll

EGSB 22%

7%

20022007, N = 610

UASB 34%

Expanded Bed Reactors: 57% Granular Sludge Bed Based: 89%

CSTR 4%

LAG 1% HYBR 2%

Figure 16.16 Implemented anaerobic technologies for industrial wastewater pictured for the period 1981-2007 (left) and the period 2002-2007 (right). UASB: upflow anaerobic sludge blanket, EGSB: expanded granular sludge bed, IC : internal circulation reactor, t ype of EGSB system with biogas-driven hydrodynamics, AF: anaerobic filter, CSTR: continuous stirred tank reactor, Lag.: anaerobic lagoon, Hybr.: combined hybrid system with sludge bed at the bottom section and a filter in top, FB: fluidized bed reactor (van Lier, 2007)

concept is the difficulty of maintaining the required contact between si udge and wastewater, because clogging of the 'bed' easily occurs. This particularly is the case for partly soluble wastewaters. These clogging problems obviously can be overcome (at least partly) by applying a primary settler and/or a pre-acidification step (Seyfried, 1988). However, this would require the construction and operation of additional units. Moreover, apart from the higher costs, it would not completely eliminate the problem of short-circuiting (clogging of the bed) flows , leading to disappointing treatment efficiencies.

Since 1981, about 140 full scale UAF installations have been put in operation for the treatment of various types of wastewater, which is about 6% of the total amount of installed high-rate reactors (Figures 16.2 and 16.15). The experiences with the system certainly are rather satisfactory, applying modest to relatively high loading rates up to I 0 kgCOD/m3.d. The UAF system will remain attractive for treatment of mainly soluble types of wastewater, particularly when the process of sludge granulation will not proceed satisfactory. On the other hand, long tenn problems related to system clogging and the stability of filter material caused a decline in the number of installed full scale AF systems. ln the last 5 years only 6 new and registered AF systems were constructed which is about 1% of the total amount of newly installed An WT systems (Figure 16.16).

ln order to mJmmise c logging and sludge accumulation in the interstices of the fi Iter material anaerobic filters are sometimes operated in a downflow mode, the so-called down-flow fixed film reactors. Various modes of operation and filter material were investigated but full-scale application 1s rather disappointing. The limiting factor is the applicable low organic loading rate owing to the limited amount of biomass that can be retained in such a system as it is primarily based on attachment of biomass to the surface of the packing material. ln UAF filters the majority of the anaerobic activity is found in the non-attached biomass.

16.7.2.3 Anaerobic Sludge Bed Reactors (ASBR) The anaerobic sludge bed reactors (ASBR) undoubted ly are by far the most popular An WT systems so far. The sludge retention in such a reactor is based on the formation of easily settling sludge aggregates (floes or granules), and on the application of an internal gas-liquid-solids separation system (GLSS device).

By far the best known example of this concept is the upflow anaerobic sludge bed reactor (UASB), which was developed in the Netherlands in the early seventies (Lettinga et al., 1976, 1980). ln view of its prospects, and the fact that almost 90% of the newly installed high-rate reactors are sludge bed systems (Figure 16. 16), the UASB process will be elaborated in more detail than the other systems (Section 16.8). At the start of 2007, about

A BIOGAS

TREATED EFFLUENT

CAS COLLECTION DOME

RISING BIOGAS

B

Granular Sludge

Blanket

437

biogas

SLUDGE BLANKET Granular I I"W;:,o;r)j'l~~:::::;;~ biomass

bed

influent influent influent influent

Figure 16.17. UASB reactors of the major anaerobic system manufacturers: (A) Paques B.V. and (B) Biothane B.V.

1,750 full-scale UASB installations have been put into operation. Most of these full scale reactors are used for treating agro-industrial wastewater, but its application for wastewater from chemical industries and sewage is increasing (Table 16.1 0). Figure 16.1 7 shows a schematic representation of a UASB reactor. Two examples of a full-scale UASB instalations are shown in figure 16. 18.

Similar to the UAF system the wastewater moves in an upward mode through the reactor. However, contrary to the AF system generally no packing material is present in the reactor vessel. The sludge bed reactor concept is based on the following ideas:

I) Anaerobic sludge has or acquires good sedimentation properties, provided mechanical mixing in the reactor remains gentle and the process is operated

correctly. For that reason, but also because it reduces the investment and maintenance costs, mechanical mixing is not applied in UASB reactors. Because of the excellent sett! ing characteristics of the sludge, high superficial liquid velocities can be applied without any risk of considerable sludge wash-out.

2) The required good contact between the sludge and wastewater in UASB-systems genera lly is accomplished (i) by feeding the wastewater as uniformly as possible over the bottom of the reactor, or (ii) as a result of the agitation caused by the production of biogas.

3) Particularly with low strength wastewater, reactors with a high height-diameter ratio are used reaching heights of 20-25 m (see section 16.7.2.4). A low surface area will facilitate the feeding of the system, whereas the accumulating biogas production over the height of the tower reactor will cause a turbulent

Figure 16.18 UASB installations for treatment of (A) f ru it ju ice f actorywastewater in Bregenz, Germany and ( B) diary w astewater in Indonesia (photo: Paques B.V. and Biothane B.V. respectively)

flow. Also the increased upflow velocity results in a better contact between the sludge and the pollutants. With wastewaters containing biodegradable additionally achieved by applying a liquid recirculation flow. As a result, a more completely mixed flow pattem is acquired and stratification of the substrate and intermediate products over the height of the reactor is minimised, thereby minimising potential inhibition.

4) The washout of sludge aggregates is prevented by separating the produced biogas using a gas collection dome installed at the top of the reactor. In this way a zone with relatively little turbulence is created in the uppermost part of the reactor, consequently the reactor is equipped with an in-built secondary clarifier. The gas collection dome acts like a three phase GLSS. The GLSS device constitutes an essential part of a UASB reactor and serves to:

a) Collect, separate and discharge the produced biogas. For a satisfactory perfom1ance the gas-liquid surface area within the device should be sufficiently large, so that gas can evade easily. This is particularly important in case scum layers should develop. Sufficient mixing by biogas turbulence should prevail at the gas-liquid interface in order to combat this phenomenon. Since the formation of scum layers is a very complex phenomenon with a wide diversity of appearance, it is impossible to give unified and clear guidelines for the dimensions of the gas-liquid interface.

b) Reduce liquid turbulences 111 the settler compartment for enhancement of sludge settling, resulting from the gas production. In order to prevent biogas bubbling to the settling zone at the top, one or more baffles should be installed beneath the aperture between the gas domes as well as between gas dome and reactor wall.

c) Remove sludge particles by a mechanism of sedimentation, flocculation and/or entrapment in a sludge blanket (if present in the settler). The collected sludge can slide back into the digestion compartment, in case the sludge bed does not reach into the settler, or can be discharged occasionally together with excess sludge from the digester compartment.

d) Limit the expansion of the sludge bed in the digester compartment. The system more or less acts as a barrier against excessive expansion of the lighter part of the sludge bed. In case the sludge bed expands into the settler, the sludge will tend to thicken (because the gas has been separated). This

le ) I i J

thickened, heavier sludge, present in the settler, lays on the top of the more voluminous sludge blanket that tends to move into the settler.

e) Reduce or prevent buoying sludge pariicles of being rinsed out from the system. For this purpose a skim layer baffle should be installed in front of the effluent weir of the overflow. Such a baffle pariicularly is essential for treating very low strength wastewaters, because wash out of viable biomass then should be kept at very low levels.

t) Aaccomplish some polishing of the wastewater with respect to suspended matter.

Some researchers and practitioners suggest replacing the GLSS device by a packed bed in the upper part of the reactor. This so-called upflow hybrid reactor is a merge between the UASB and the UAF reactors. ln some designs the packing material is mounted only in the settling compartment leaving the GLSS at its original position. About 2 to 3% of all anaerobic reactors installed are hybrid reactors (see Figure 16.16). In most applications, the majority of organic matter conversion is located in the sludge bed section whereas the removal of a specific fraction of pollutants is located in the filter area at the top. Specific chemical wastewaters show better treatment efficiencies for all compounds using hybrid systems compared to UASB reactor. The most known example is the treatment of purified therephthalic acid (PT A) wastewater (Kleerebezem, l999a,b ). Results showed that the conversion of therephthalic acid to benzoate is only possible at low concentrations of acetate and benzoate. By applying a hybrid system, the latter two are converted in the sludge bed area whereas, therephthalic acid is then converted in the hybrid section, where specific flora is retained for degrading the refractory compound. The most known disadvantage of hybrid reactors is the deterioration of the filter section after prolonged periods of operation. Hybrid reactors are also advantageous for achieving enhanced effluent polishing as colloidal matter is entrapped at the top part of the system. In fact, trials with domestic sewage showed improved removal of both suspended solids and colloidal matter (Elmitwalli et al., 2002). Biomass accumulating in the packing material ensures a prolonged contact of wastewater with viable bacterial matter, in the absence of packing material little viable biomass will be present in the upper part of the reactor due to the sludge discharge regime generally applied in anaerobic sewage treatment plants. The packing material furthermore enhances flocculation of the finer suspended solids fraction present in the wastewater.

,T]

16.7.2-4 Anaerobic expanded and fluidized bed systems (EGSB and FB)

Expanded bed and fluidized bed systems are regarded as the second generation of sludge bed reactors achieving extreme organic loading rates (exceeding 30 to 40 kgCOD/m3 d) . The FB process is based on the occurrence of bacterial attachment to mobile carrier particles, which consist, for example, of fine sand (0.1-0.3 mm), basalt, pumice, or plastic. The FB system can be regarded as an advanced anaerobic technology (Li and Sutton, 1981; Heijnen, 1983, 1988), that may reach loading rates of 50-60 kgCOD/m3 d. However, long-term stable operation appears to be problematic. The system relies on the fom1ation of a more or less uniform (in thickness, density, strength) attached biofilm and/or particles. ln order to maintain a stab le situation with respect to the biofilm deve lopment, a high degree of pre-acidification is considered necessary and dispersed mat